Hydrol. Earth Syst. Sci., 20, 1–15, 2016
www.hydrol-earth-syst-sci.net/20/1/2016/
doi:10.5194/hess-20-1-2016
© Author(s) 2016. CC Attribution 3.0 License.
Carbon and nitrogen dynamics and greenhouse gas emissions in
constructed wetlands treating wastewater: a review
M. M. R. Jahangir1,2 , K. G. Richards2 , M. G. Healy3 , L. Gill1 , C. Müller4,5 , P. Johnston1 , and O. Fenton2
1 Department
of Civil, Structural & Environmental Engineering, Trinity College Dublin, Dublin 2, Ireland
of Environment, Soils & Land Use, Teagasc Environment Research Centre, Johnstown Castle,
Co. Wexford, Ireland
3 Civil Engineering, National University of Ireland, Galway, Co. Galway, Ireland
4 School of Biology and Environmental Science, University College Dublin, Belfield, Dublin, Ireland
5 Department of Plant Ecology (IFZ), Justus-Liebig University Giessen, Giessen, Germany
2 Department
Correspondence to: M. M. R. Jahangir (jahangim@tcd.ie)
Received: 19 June 2014 – Published in Hydrol. Earth Syst. Sci. Discuss.: 4 July 2014
Revised: 28 October 2015 – Accepted: 29 November 2015 – Published:
Abstract. The removal efficiency of carbon (C) and nitrogen
(N) in constructed wetlands (CWs) is very inconsistent and
frequently does not reveal whether the removal processes are
due to physical attenuation or whether the different species
have been transformed to other reactive forms. Previous research on nutrient removal in CWs did not consider the dynamics of pollution swapping (the increase of one pollutant
as a result of a measure introduced to reduce a different pollutant) driven by transformational processes within and around
the system. This paper aims to address this knowledge gap by
reviewing the biogeochemical dynamics and fate of C and N
in CWs and their potential impact on the environment, and by
presenting novel ways in which these knowledge gaps may
be eliminated. Nutrient removal in CWs varies with the type
of CW, vegetation, climate, season, geographical region, and
management practices. Horizontal flow CWs tend to have
good nitrate (NO−
3 ) removal, as they provide good conditions
for denitrification, but cannot remove ammonium (NH+
4 ) due
+
to limited ability to nitrify NH4 . Vertical flow CWs have
good NH+
4 removal, but their denitrification ability is low.
Surface flow CWs decrease nitrous oxide (N2 O) emissions
but increase methane (CH4 ) emissions; subsurface flow CWs
increase N2 O and carbon dioxide (CO2 ) emissions, but decrease CH4 emissions. Mixed species of vegetation perform
better than monocultures in increasing C and N removal and
decreasing greenhouse gas (GHG) emissions, but empirical
evidence is still scarce. Lower hydraulic loadings with higher
hydraulic retention times enhance nutrient removal, but more
empirical evidence is required to determine an optimum design. A conceptual model highlighting the current state of
knowledge is presented and experimental work that should
be undertaken to address knowledge gaps across CWs, vegetation and wastewater types, hydraulic loading rates and
regimes, and retention times, is suggested. We recommend
that further research on process-based C and N removal and
on the balancing of end products into reactive and benign
forms is critical to the assessment of the environmental performance of CWs.
1
Introduction
Increasing anthropogenic loading of reactive nitrogen (Nr;
all forms of nitrogen (N) except di-nitrogen gas, N2 ) along
the N cascade in the environment raises many critical concerns for human health, drinking water quality (Gray, 2008),
coastal and marine water degradation as well as algal blooms
and hypoxia (Conley et al., 2009; Rabalais et al., 2010). Constructed wetlands (CWs) are artificial sinks for Nr (Galloway
et al, 2003; Tanner et al., 2005), and have been successfully
used to treat domestic sewage, urban runoff and storm water,
industrial and agricultural wastewater, and leachate. While
the biogeochemistry of wetlands in general has been discussed in the literature (Whalen, 2005; Reddy and Delaune,
2008), less is known about the delivery pathways of the transformation products of carbon (C) and N from CWs treating
Published by Copernicus Publications on behalf of the European Geosciences Union.
2
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
wastewater. Although CWs have a proven potential for organic C and N removal, with few exceptions (Dzakpasu et
al., 2014), studies have rarely quantified all relevant pathways. This has meant that reported removal efficiencies have
been variable (Seitzinger et al., 2002). If the fate of C and N
is accurately quantified, appropriate design and management
strategies may be adopted.
Constructed wetlands are complex bioreactors that facilitate a number of physical, chemical, and biological processes, but are frequently evaluated as a black box in terms
of process understanding (Langergraber, 2008). Many investigations target single contaminant remediation and disregard
the reality of mixed contaminants entering and leaving CWs.
They do not consider the dynamics of pollution swapping
(the increase in one pollutant as a result of a measure introduced to reduce a different pollutant) driven by transformational processes within and around the system. This means
that potential negative impacts that CWs may have on the environment, such as greenhouse gas (GHG) emissions (IPCC,
2013; Clair et al., 2002; Mander et al., 2008; Mitsch and Gosselink, 2000) or enhancement of pollution swapping (Reay,
2004), are not accounted for in analyses. There are many
pathways by which the removed N can contribute to water and air pollution: accumulation and adsorption in soils,
+
leaching of nitrate (NO−
3 ) and ammonium (NH4 ) to groundwater, emissions of nitrous oxide (N2 O) and ammonia (NH+
3)
to the atmosphere, and/or conversion to N2 gas. Constructed
wetlands significantly contribute to atmospheric N2 O emissions either directly to the atmosphere from the surface of
the wetland (IPCC, 2013; Søvik et al., 2006; Ström et al.,
2007; Elberling et al., 2011) or indirectly via dissolved N2 O
in the effluent or groundwater upon discharge to surface waters. The IPCC (2013) has recognized the significance of indirect N2 O emissions from CW effluent that is discharged
to aquatic environments, and estimate emission factors (EF)
ranging from 0.0005 to 0.25. Production and reduction processes of N2 O in the environment are not yet fully understood.
Constructed wetlands receive organic C from the influent wastewater and from fixation by the photosynthetic hydrophytes, which are incorporated into soil as organic C. Soil
organic C undergoes the biogeochemical processes that regulate C accretion in soil and microbial respiration, producing
carbon dioxide (CO2 ). Anaerobic mineralization of organic
C by methanogenic archaea can produce methane (CH4 )
(Laanbroek, 2010; Ström et al., 2007; Søvik et al., 2006; Pangala et al., 2010). Constructed wetlands can also contribute to
the dissolved organic carbon (DOC) load transfer to groundand surface waters, which may produce and exchange substantial amounts of CO2 and CH4 with the atmosphere (Clair
et al., 2002; Elberling et al., 2011). Therefore, CWs can diminish the environmental benefits of wastewater treatment.
The dynamics of dissolved N2 O, CO2 , and CH4 in CWs is a
key knowledge gap in global GHG budgets.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
Surface emissions of GHG from CWs have been commonly measured by the closed chamber method (Johansson
et al., 2003, 2004; Mander et al., 2005, 2008), but have rarely
been measured by ebullition and diffusion methods (Søvik
et al., 2006). The measured rates have shown high spatial,
temporal, and diurnal variations due to the change in biogeochemistry of C and N and plant–microbe–soil interaction
over time and space. Surface emissions cannot explain the kinetics of production and consumption rates of GHG, which
we need to know in order to adopt better management practices to mitigate emissions. In addition, subsurface export of
dissolved nutrients and GHG, an important pathway of nutrient loss (Riya et al., 2010), is frequently ignored. Mass
balance analysis of the different components of the N cycle and kinetics of their transformation processes occurring
within the treatment cells using the isotope-tracing 15 N technique can provide mechanistic information for N transformation products (Lee et al., 2009; O’Luanaigh et al., 2010)
and may be used to start to answer such questions. Similarly, 14 C application and measurement of C species (e.g.
CO2 , CH4 , and DOC) may elucidate the C mineralization and
CO2 and CH4 production and consumption. Used in combination, these methods may provide a comparative analysis of
the rates of C and N transformation processes and the role
+
of these processes in delivering NO−
3 , NH4 , and DOC to
ground/surface waters and N2 O, CO2 , and CH4 to the atmosphere.
Past reviews on CWs, though very limited, summarize the
performance of different types of CWs on C and N removal
(Vymazal, 2007) and surface emissions of GHG (Mander et
al., 2014), but have not discussed the mechanisms of nutrient removal and the fate of the nutrients delivered and removed to and from CWs. Therefore, the objectives of this
review are to (i) understand the biogeochemical dynamics of
C and N in CWs, (ii) better understand the fate of various C
and N species in a holistic manner, in addition to the conventional influent/effluent balance for nutrient removal, (iii)
identify the research gaps that need to be addressed to optimize nutrient removal and mitigate GHG emissions, and (iv)
discuss emerging measurement techniques that may give insights into the production and reduction of GHG.
2
Removal efficiency, hydraulic loading, and retention
time
In CWs, the efficiency of C and N removal is generally limited and highly variable over CW types, plant types, seasons, climatic regions, and management practices. On average, it appears that 50 and 56 % of the influent total nitrogen
(TN) and total organic carbon (TOC), respectively, can be
removed, but the removal rates are very inconsistent. Mean
(± standard error) TN removals, obtained from the literature
cited in this paper, ranged from 31.3 ± 6.3 % in surface flow
(SF) CWs to 40.4 ± 4.4 % in subsurface flow (SSF) CWs,
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M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
3
Table 1. TN input (mg N L−1 ), TN output (TN, mg N L−1 ), and TN removal (%) in various CWs treating wastewater; average standard error
(± SE) is presented for TN removal; NA – data not available.
CW type
N input (mg N L−1 )
Treatment
N output (mg N L−1 )
N removal ( %)
References
TN
NH+
4
NO−
3
TN
NH+
4
NO−
3
TN
NH+
4
NO−
3
0.01 ± 3.0
61.2 ± 1.7
32.6 ± 1.9
NA
NA
NA
NA
NA
NA
NA
NA
0.02 ± 6.7
0.3 ± 0.09
0.9 ± 0.4
NA
NA
NA
NA
NA
NA
NA
NA
21.4
2.1
15.4
61
97–98
39–48
39
35
> 90
45
40
31.3 ± 6.3
66.7
3.6
21.4
NA
NA
NA
NA
NA
NA
NA
NA
93.3
57.1
−800
NA
NA
NA
NA
NA
NA
NA
NA
Søvik et al. (2006)
Søvik et al. (2006)
Søvik et al. (2006)
Song et al. (2011)
Dzakpasu et al. (2011)
Vymazal (2007)
Vymazal (2010)
Shamir et al. (2001)
Harrington et al. (2007)
Toet et al. (2005)
Van der Zaag et al. (2010)
52.1
15.7
42.2
29
46–48
28
78
47
20
30
40
40.4 ± 4.4
56.9
−12.2
NA
NA
−2850
92.9
NA
NA
50
39
20
33
30
NA
−29
8
31
33
Søvik et al. (2006)
Søvik et al. (2006)
Van der Zaag et al. (2010)
O’Luanaigh et al. (2010)
Mander et al. (2008)
Van der Zaag et al. (2010)
Kato et al. (2006)
Vymazal and Kröpfelova (2010)
Vymazal and Kröpfelova (2010)
Vymazal and Kröpfelova (2010)
Vymazal and Kröpfelova (2010)
15.3
9.1
49.3 ± 1.8
74 ± 3
37.0 ± 10.9
11.2
56.9
NA
NA
−54.5
−25 400
NA
NA
Søvik et al. (2006)
Søvik et al. (2006)
Yan et al. (2012)
Zhao et al. (2014)
SF_Finland
SF_Finland
SF_Norway
SF
SF
SF
SF
SF
SF
SF
SF
All SF
Municipal
Agril. runoff
Municipal
Municipal
Domestic
Various
Municipal
Municipal
various
Municipal
Dairy washout
1.4 ± 150
66.1 ± 1.9
43.4 ± 3.6
NA
NA
NA
NA
NA
NA
NA
227
0.03 ± 5.8
63.5 ± 1.3
41.5 ± 3.0
4.5
40
39
36
196
80
0.95
NA
0.3 ± 95
0.7 ± 0.13
0.0 ± 0.0
15.5
5
4.4
<2
<1
1.54
NA
1.1 ± 48
64.7 ± 1.7
36.7 ± 2.7
NA
NA
NA
NA
NA
NA
NA
NA
HSSF_Estonia
HSSF_Norway
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
All HSSF
Municipal
Municipal
Dairy washout
Domestic
Domestic
Dairy washout
Milk parlour
Agriculture
Industry
Landfill
Municipal
96.5 ± 3.0
53.4 ± 4.3
306 ± 101*
NA
87
227
112
67
124
157
43
83.9 ± 2.7
38.4 ± 7.7
NA
74.9
0.2 ± 0.02
14.1 ± 7.5
NA
3.9
46.2 ± 1.5
45.0 ± 4.1
177 ± 58*
NA
36.2 ± 1.4
43.1 ± 4.7
NA
NA
5.9 ± 0.65
1.0 ± 0.8
NA
NA
22
40
65
149
24
NA
0.85
8.5
1.5
2
24
27
103
147
24
11
11
31
98
14
NA
1.1
7.4
1.3
1.2
VSSF_Estonia
VSSF_Norway
VSSF
VSSF
All VSSF
Municipal
Municipal
Municipal
Municipal
50.9 ± 9.2
52.6 ± 5.2
41.0 ± 0.5
46 ± 13
35.7 ± 6.2
49.6 ± 4.0
NA
NA
1.1 ± 0.32
0.0 ± 0.0
NA
NA
43.1 ± 7.6
47.8 ± 6.9
20.7 ± 0.8
NA
31.7 ± 5.5
21.4 ± 6.9
NA
NA
1.7 ± 0.84
25.5 ± 1.3
NA
NA
SF – surface flow; HSSF – horizontal subsurface flow; VSSF – vertical subsurface flow; ∗ mg N m−2 h−1 .
Table 2. Total organic C (TOC) removal (%) in various CWs treating wastewater; average standard error (± SE) is presented for TOC
removal; NA – data not available.
CWs type
Treatment
C input (TOC; mg C L−1 )
C output (TOC; mg C L−1 )
TOC Removal (%)
References
SF_Finland
SF_Finland
SF_Norway
SF
All SF
Municipal
Agril runoff
Municipal
Dairy wash out
13.0 ± 0.3
25.0 ± 3.4
26.7 ± 2.9
186a
14.0 ± 0.5
20.0 ± 3.4
17.1 ± 1.8
136a
−7.7
20.0
36.0
27
18.8 ± 9.4
Søvik et al. (2006)
Søvik et al. (2006)
Søvik et al. (2006)
Van der Zaag et al. (2010)
HSSF
HSSF
HSSF_Estonia
HSSF_Norway
All HSSF
Domestic
Dairy wash out
Municipal
Municipal
150b
186a
62.8 ± 16.6a
40.5 ± 11.3
NA
107.9a
41.0 ± 11.3a
15.0 ± 2.4
NA
42
34.7
63.0
46.6 ± 7.3
Garcia et al. (2007)
Van der Zaag et al. (2010)
Søvik et al. (2006)
Søvik et al. (2006)
VSSF_Estonia
VSSF_Norway
VSSF
VSSF
All VSSF
Municipal
Municipal
Municipal
Municipal
132.2 ± 32.2a
40.5 ± 11.3
106 ± 35
249 ± 49
62.8 ± 16.6 a
15.0 ± 2.4
74 ± 21
NA
52.5
63.0
26 ± 4.6
83 ± 1.0
56.2 ± 9.5
Søvik et al. (2006)
Søvik et al. (2006)
Yan et al. (2012)
Zhao et al. (2014)
SF – surface flow; HSSF – horizontal subsurface flow; VSSF – vertical subsurface flow; a BOD; b mg m−2 h−1 .
whereas TOC removal ranged from 18.8 ± 9.4 % in SF CWs
to 56.2 ± 9.5 % in vertical subsurface flow CWs (Tables 1
and 2). In European systems, for example, typical removals
of ammoniacal N in long-term operation are around 35 %,
but can be enhanced if some pre-treatment procedures are
followed (Verhoeven and Meuleman, 1999; Luederitz et al.,
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2001). Generally, TN removal is higher in SF CWs than SSF
CWs (Table 1), but studies differ. For example, Van der Zaag
et al. (2010) showed higher N removal in SF CWs than SSF
CWs, but Søvik et al. (2006) and Gui et al. (2007) showed the
opposite. In SSF CWs, limited removal can be caused by a
reduced environment that enhances NH+
4 accumulation and
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
4
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
limits NH+
4 oxidation. In SF CWs, denitrification rates can
be limited due to lack of NO−
3 . In vertical subsurface flow
−
(VSSF) CWs, aeration can increase NH+
4 oxidation to NO3 ,
+
which can be denitrified or converted to NH4 by dissimila+
tory NO−
3 reduction to NH4 (DNRA).
Plant species are important components of CWs, and affect C and N removals. Optimal species selection for best
removal is difficult because some species are efficient in removing one pollutant but not the other (Bachand and Horne,
2000; Bojcevska and Tonderski, 2007; da Motta Marques et
al., 2001). In some studies there are no inter-species differences at all (Calheiros et al., 2007). Mixed species perform
better than monocultures to remove C and N pollutants because they increase microbial biomass and diversity. Payne
et al. (2014a) discussed the role of plants in nutrient removal.
Plants regulate CW hydrology (evaporation and transpiration) and temperature (insulating CWs from seasonal temperature change, trapping falling and drifting snow, and heat
loss of wind). Some species can create a large surface area
for microbial attachment and enhance microbial diversity, but
experimental evidence is still scarce.
Soil physico-chemical properties, such as permeability
(Dzakpasu et al., 2014) and cation exchange capacity (Drizo
et al., 1999) are important factors controlling the purification
capacity in CWs. Microbial activities and growth depend on
substrate C quality and C : N ratios, which affect nutrient removal. Growth of heterotrophic microorganisms is a function
of the wastewater C : N (Makino et al., 2003). High C : N ratios can enhance denitrification by providing electron donors
for denitrifiers, but the opposite can increase nitrification.
High C : N ratios can also encourage DNRA over denitrification. Yan et al. (2012) measured a high TN removal but low
TOC removals at a C : N ratio 2.5 : 1, which indicates that
removal of one parameter might lead to a problem with a different one. The uncertainty in the conditions for achievement
of optimum removal suggests that the rates of C and N transformations and the fate of the removed nutrients within CWs
should be investigated. However, to our knowledge, no study
has provided a holistic evaluation of C and N attenuation and
transformation.
The removal of pollutants in CWs depends on hydraulic
loading rates (HLR) and hydraulic retention time (HRT)
(Toet et al., 2005). The HLR and HRT are considered to
be significant design parameters determining the nutrient removal efficiencies (Weerakoon et al., 2013). Longer HRTs of
wastewater in CWs increase the removal of C and N (Wang
et al., 2014) by increasing sedimentation and duration of contact between nutrients and the CWs. The effects of HLR and
HRT can vary with the nature of the use of CWs, e.g. whether
they are used for treating single or mixed pollutants. To reduce Nr delivery to the receiving waters or to the atmosphere,
CWs need to be optimally designed with respect to HLR and
HRT.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
Fluctuating hydraulic loading influences all biotic and abiotic processes in CWs (Mander et al., 2011). For example,
if the groundwater table is lowered through changes in hydraulic loading, soil aeration can increase or decrease. Ammonification and nitrification rates increase with increased
soil aeration and this enhances C utilization by bacteria and,
therefore, can stimulate the removal of C and N. Investigation
into the effects of fluctuating hydraulic loadings (hydraulic
pulsing) on C and N removals and their transformation products will provide information about the fate of the added nutrients in terms of their environmental benefits and/or pollution swapping potential. For example, if the dominant product is N2 , the system will be relatively benign in terms of
its impact on the environment, but if it is NH+
4 , it can be
fixed in the soils or transported to ground- and surface waters
connected to CWs if the cation exchange sites become saturated. Several authors have used a wide range of HLRs and
HRTs to measure nutrient removal efficiency, but experimental evidence linking HLR and HRT to removal efficiency is
scarce (Toet et al., 2005). Luo et al. (2005) reported that low
HLR results in incomplete denitrification, whereas Zhang et
al. (2006) argued that low HLR increases NH+
4 and chemical oxygen demand oxidation. The way in which the performance of a CW is assessed can lead to different conclusions
regarding the removal of Nr. For future studies, evaluation of
CWs in a holistic manner, which includes pollution swapping
at different HLRs and HRTs, is important, particularly within
the context of the changing hydrologic cycle in a changing
climate. In addition, local legislative targets should be considered and weighting factors (e.g. the relative importance
of, say, GHG over water quality targets) should be developed
to evaluate the overall performance of CWs. In addition to
the estimation of nutrient removal rates, investigation of the
effect of HLR and HRT on the different forms of nutrients
in the final effluent and their fate in the natural environment
may help elucidate the pollution swapping potential of CWs.
3
Accumulation of C and N in CWs soils
The soil in CWs is a major sink for C and N (Mustafa and
Scholz, 2011). However, although data on the influent and
effluent N concentrations are available, data on N accumula−
tion (dissolved organic nitrogen (DON), TN, NH+
4 , or NO3 –
N) within the soil profile of various CWs are scarce. The
wide range of N accumulation reported in the literature (e.g.
30–40 %, Shamir et al., 2001; 39 %, Harrington et al., 2007;
9 %, Mander et al., 2008; 2.5 %, Obarska-Pempkowiak and
Gajewska, 2003) may be due to the variations in CW types
and management strategies. The accumulated species of N
are reactive unless they have been transformed to N2 by biogeochemical processes. However, there is a dearth of information on the extent of Nr accumulation in soils and discharge to surface waters and air (Shamir et al., 2001). Accu−
mulated organic N could be mineralized to NH+
4 and NO3 ,
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M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
depending on the physico-chemical properties of soil. The
Nr could be assimilated by plants and microbes, which are
recycled in a soil–plant–soil continuum. Nitrogen spiralling
+
occurs from NH+
4 to organic N and back to NH4 within
the CW (O’Luanaigh et al., 2010). Typically, N accumulation has been found to decrease with soil depth (Shamir et
al., 2001). In terms of the conventional input–output balance,
these are considered as removed N, but may, in fact, remain
in such a biogeochemically active system. In addition to N,
organic C accumulation occurs in CW soils (Nguyen, 2000).
Soils of CWs represent organic C and Nr-rich systems,
where the products of continuously occurring biogeochemical processes, such as accumulation in soil and transportation
to fresh waters and to the atmosphere, need to be quantified.
Such an approach will show the shortcomings of conventional removal efficiency estimation methods and will also
demonstrate how the apparently removed C and N species
can become a source of contamination. Estimation of the
rates of nutrient accumulation in soils in various types of
CWs under different management systems is important. The
stability of the accumulated C and N under changing climatic
scenarios also needs to be addressed to consider the longterm sustainability of CWs.
4
C and N dynamics and greenhouse gas emissions
Increased nutrient input to CWs increases the productivity of
wetland ecosystems and the production of GHG. As CWs are
designed to remove pollutants in an anaerobic/suboxic environment, they change the C and N biogeochemistry and contribute significantly to CH4 and N2 O emissions (Johansson,
2002, Johansson et al., 2003; Mander et al., 2005, 2008; Stadmark and Leonardson, 2005; Liikanen et al., 2006). Søvic et
al. (2006) measured N2 O, CH4 , and CO2 emissions in various CWs in different European countries, and suggested that
the potential atmospheric impacts of CWs should be examined as their development is increasing globally. Management of CWs must consider the negative climatic aspects
of increased emissions of GHG in addition to their primary
functions (Ström et al., 2007). Therefore, estimation of the
contribution of CWs to global warming is required. In this
regard, measurement of spatial and temporal variations (seasonal and diurnal) of GHG emissions is necessary to accurately estimate CW-derived GHG emissions. A holistic assessment of ecologically engineered systems has been outlined by Healy et al. (2011, 2014) and developed further by
Fenton et al. (2014). Such assessments can be applied in evaluating nutrient dynamics in CWs. Moreover, plant mediated
GHG emissions could be an important component of total
emissions, but again research in this area is very limited. Effective modelling or up-scaling of GHG emissions from watershed to regional/national scales is important for the improvement of global GHG budgets. Such up-scaling needs
an accurate estimation of C and N inputs and outputs, i.e. a
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5
balance coupled with net GHG emissions, while considering
all possible processes and pathways involved. A study of the
dynamics of C and N in CWs is crucial, as the forms of removed C and N are particularly pertinent to their potential for
pollution swapping, global warming, and water pollution.
Processes involved in N removal and N transformations
in wetlands include sedimentation of particulates (Koskiaho, 2003), nitrification, denitrification, DNRA (Poach et al.,
2003; Burgin et al., 2013), microbial assimilation and plant
uptake–release (Findlay et al., 2003), and anammox (anaerobic ammonium oxidation) and deamox (DEnitrifying AMmonium OXidation). Constructed wetlands are complex systems that facilitate aerobic and anaerobic microsites (Wynn
and Liehr, 2001). Nitrification, denitrification, and nitrifier
denitrification are the processes responsible for the production of N2 O. Depending on the environmental conditions or
management practices prevailing, a certain process will dominate; e.g., denitrification is the dominant process in SF CWs
(Beaulieu et al., 2011), but nitrifier denitrification is dominant in VSSF CWs (Wunderlin et al., 2013). Generally, CWs
are anaerobic but aquatic macrophytes can transport oxygen from the atmosphere to the rooting zone, where it can
sustain nitrification. The existence of microsites that support
high activity and promote denitrification has been shown in
soils (Parkin, 1987). Such conditions are also likely to occur
in CWs, which have patchy distributions of organic material (e.g. particulate organic carbon), due to rhizodepositions
(Minett et al., 2013; Hamersley and Howes, 2002). Minett
et al. (2013) found that simultaneous oxygenation of the rhizosphere, through radial oxygen loss, and enhanced oxygen
consumption by the soil occurs in the area immediately surrounding the roots. Nitrate produced in the rooting zone can
be taken up by plants or denitrified and/or converted back to
NH+
4 by DNRA.
Competition for NO−
3 may occur between denitrification
and biotic assimilation. This is likely governed by the prevailing aerobic/anaerobic conditions and therefore dependent
on the type of wetland. For instance, in storm water biofiltration systems, prolonged periods of inundation and dry periods may support bio-assimilation over denitrificaton (Payne
et al., 2014a, b).
The conditions that favour the occurrence of either denitrification or DNRA are still in debate (Rütting et al., 2011).
DNRA is thought to be favoured by a C : NO−
3 ratio of > 12
(Rütting et al., 2011) and occurs at low levels of oxidationreduction potential (Thayalakumaran et al., 2008). The differences between denitrification and DNRA may be due to
the availability of organic matter, because DNRA is favoured
at a high C : NO−
3 ratio and denitrification is favoured when
carbon supplies are limiting (Korom, 2002; Kelso et al.,
1997). The fermentative bacteria that carry out DNRA are
obligate anaerobes, and so cannot occupy all the niches that
denitrifiers can (Buss et al., 2005). Takaya (2002) stated that
a more reducing state favours DNRA over denitrification.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
6
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
Pett-Ridge et al. (2006) showed that DNRA is less sensitive
to dissolved oxygen (DO) than denitrification. Fazzolari et
al. (1998) showed that the partitioning between DNRA and
denitrification depends on the C : NO−
3 ratio and C rather than
DO.. Significant DNRA may occur only at a C : NO−
3 ratio
above 12 (Yin et al., 1998). Different numbers of electrons
are required in the reduction of each NO−
3 molecule: five
for denitrification and eight for DNRA. Therefore, more organic matter can be oxidized for each molecule of NO−
3 by
DNRA than by denitrification. In addition, NO−
reduction
3
is generally performed by fermentative bacteria that are not
dependent on the presence of NO−
3 for growth under anaerobic conditions. Therefore, DNRA bacteria may be favoured
in NO−
3 -limited conditions (Laanbroek, 1990). Recent studies have suggested that DNRA may be an important process
compared to denitrification in wetland sediments (Burgin and
Hamilton, 2008). Van Oostrom and Russell (1994) found a
5 % contribution of DNRA to NO−
3 removal in CWs. Little is
known about the eventual fate of the NO−
3 that is converted to
+
NH4 via DNRA pathways. In recent years, N-cycling studies have increasingly investigated DNRA in various ecosystems to explore its importance in N cycling (Rütting et al.,
2011), but controls on DNRA are relatively unknown (Burgin
et al., 2013), DNRA being probably the least studied process
of N transformation in wetlands (Vymazal, 2007). However,
DNRA can be a significant pathway of NO−
3 reduction that
impacts on the CW ecosystem services and so should therefore be evaluated.
Denitrification has been estimated to be a significant N
removal process, but actual quantification data are scarce.
Few studies have estimated N losses by denitrification, e.g.
19 % (Mander et al., 2008) and 86 % (Obarska-Pempkowiak
and Gajewska, 2003) of the total N input based on the mass
balance study. To our knowledge, no data are available on
denitrification measurements in soil/subsoils of surface flow
CWs. While many of these pathways transfer Nr (mainly
NH+
4 and N2 O) to the environment, other pathways can convert Nr to N2 (e.g. denitrification, anammox, and deamox).
+
Anammox can remove NO−
2 and NH4 as N2 when the existing environment is hypoxic. Deamox can remove NO−
3 and
−
−
NH+
as
N
,
where
NO
is
converted
to
NO
by
autotrophic
2
4
3
2
denitrification with sulfide (Kalyuzhnyi et al., 2006). In CWs,
anammox and deamox are not well understood, so it is crucial to identify which of the processes are occurring in a specific type of CW and the rate at which they occur. Once a
process that provides N2 as the end product is determined,
then the management of the CW can be directed towards enhancement of that process. Hence, quantifying the rates of
these processes for various types of CW is required for improved N management towards lowering Nr in the environment.
The various components of the C cycle include: fixation
of C by photosysnthesis, respiration, fermentation, methanogenesis ,and CH4 oxidation with reduction of sulfur, iron, and
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
NO−
3 . Anaerobic methane oxidation coupled with denitrification, a recently proposed pathway of the C cycle (á Norði
and Thamdrup, 2014; Haroon et al., 2013; Islas-Lima et al.,
2004), can reduce CH4 emissions in CWs. The C removal
processes are sedimentation, microbial assimilation, gaseous
emissions, dissolved C losses through water to ground- and
surface water bodies, and chemical fixation (bonding with
chemical ions). Net primary productivity of wetland hydrophytes varies across CW type, season, climatic region,
and local environmental conditions. For example, results can
vary remarkably for CWs containing the same plant species
in different geographical regions (Brix et al., 2001). The rate
of carbon mineralization in CW sediments depends on the
redox chemistry of soil, the bio-availability of organic C
and temperature. In particular, areas of sediment subjected
to prolonged low redox conditions (e.g, −150 mV) are conducive to methanogens and rates of CH4 emissions exceeding
132 mg m−2 d−1 (Brix et al., 2001), but this is highly variable
depending on C : N ratio of the influent water and wetland
seasonality. In summer, oxygen diffusion to the topsoil can
reduce methanogenesis and stimulate CH4 oxidation (Grünfeld and Brix, 1999). However, an increase in temperature
can decrease DO in deeper subsoil layers, which can enhance
CH4 production. As in all biochemical reactions, temperature
increases C and N turnover in CWs, causing high variations
in GHG emissions in different types of CWs in different regions (temperate/tropical/arctic). This warrants the acquisition of more measurement data across CW types and regions
for the better extrapolation of GHG emissions. The C : N ratios of wastewater affect microbial growth and development
that, in turn, affect their response to C and N cycles and GHG
emissions. Previous research on the effects of C : N ratios on
nutrient removal and GHG emissions is limited. A few examples include Yan et al. (2012) and Zhao et al. (2014), who
measured lower CO2 and CH4 emissions at C : N ratios of
between 2.5 : 1 and 5 : 1, but this lower range of C : N ratios
decreased TOC removal. Hence, investigation of the influence of C : N ratio on nutrient removal efficiencies and GHG
emissions across CW and management types is crucial.
Emissions of GHG in CWs can vary across CW typologies, e.g. surface flow or subsurface flow (Tables 3 and 4).
Generally, CH4 emissions are higher in SF CWs than in SSF
CWs (Table 3), but may vary with season, which requires
investigation. Nitrous oxide and CO2 emissions are higher
in VSSF CWs than horizontal subsurface flow (HSSF) and
SF CWs. The N2 O EF (N2 O / TN input × 100) ranged from
0.61 ± 0.21 % in SF CWs to 1.01 ± 0.48 % in VSSF CWs.
The EF for CH4 emissions ranged from 1.27 ± 0.31 % in
VSSF CWs to 16.8 ± 3.8 % in SF CWs. The GHG from CWs
can vary between vegetated and non-vegetated systems (Table 5).
Aquatic plants play an important role in GHG production
and transport to the atmosphere by releasing GHG through
their interconnected internal gas lacunas (Laanbroek, 2010).
Emergent plants can transport atmospheric oxygen to the
www.hydrol-earth-syst-sci.net/20/1/2016/
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
7
Table 3. Nitrous oxide (N2 O) emissions (mg N m−2 d−1 ); N2 emissions (mg N m−2 d−1 ) and N2 O emission factor (N2 O / TN input × 100)
in various type of CWs; mean standard error (± SE) was presented for N2 O emission factor; NA – data not available.
CW type
Treatment
Denitrification
N2 O emissions
(mg N m−2 d−1 )
N2 emissions
(mg N m−2 d−1 )
N2 O-N / TN (%)
N2 -N / TN (%)
References
HSF
HSF
HSF
HSF
HSF_Finland
HSF_Finland
HSF_Norway
All SF
Agril. tile drainage
Treated municipal
Agril. drainage
Dairy wash out
Municipal
Agril. runoff
Municipal
0.01–0.12
2.0 ± 3.3
−0.2–1.9
16.8 ± 7.0
0.01 ± 0.01
0.40 ± 0.25
4.0 ± 1.6
2.78 ± 1.72
NA
NA
NA
NA
NA
NA
NA
0.19–1.4
0.02–0.27
−0.14–0.52
0.33 ± 0.12
1.6 ± 1.3
0.37 ± 0.18
1.5 ± 4.4
0.61 ± 0.21
NA
NA
NA
NA
NA
NA
NA
Xue et al. (1999)
Johansson et al. (2003)
Wild et al. (2002)
Van der Zaag et al. (2010)
Søvik et al. (2006)
Søvik et al. (2006)
Søvik et al. (2006)
HSSF
HSSF_Estonia
HSSF_Norway
HSSF
HSSF
HSSF
HSSF
HSSF
VSSF
All HSSF
Domestic
Municipal
Municipal
Domestic
Domestic
Domestic
Dairy wash out
Domestic
Domestic
0.2–17.0
7.1 ± 1.2
6.9 ± 4.3
1.3–1.4
0.003–0.001
0.13
9.5 ± 1.5
0.17
0.17
4.23 ± 1.87
NA
NA
NA
160–170
0.01–5.42
NA
NA
NA
NA
0.06–3.8
0.05 ± 0.31
0.24 ± 0.53
0.37–0.60
NA
0.008
0.18 ± 0.12
0.23
0.01
0.62 ± 0.38
NA
NA
NA
15.2–22.7
NA
NA
NA
NA
Mander et al. (2005)
Søvik et al. (2006)
Søvik et al. (2006)
Mander et al. (2008)
Teiter and Mander (2005)
Fey et al. (1999)
Van der Zaag et al. (2010)
Liu et al. (2009)
Mander et al. (2011)
VSSF
VSSF
VSSF
VSSF
VSSF
VSSF
VSSF_Estonia
VSSF_Norway
All VSSF
Domestic
Domestic
Domestic
Domestic
Domestic
Domestic
Municipal
Municipal
0.001–0.002
4.6
11.0
1.44
0.005
0.003
15 ± 3.9
960 ± 40
123.8 ± 106
0.01–5.0
150
NA
NA
NA
NA
NA
NA
NA
0.45–0.50
0.29
0.03
0.09
0.04
04.3 ± 0.95
1.4 ± 0.72
1.01 ± 0.48
NA
NA
NA
Teiter and Mander (2005)
Mander et al. (2008)
Mander et al. (2005)
Mander et al. (2011)
Gui et al. (2007)
Liu et al. (2009)
Søvik et al. (2006)
Søvik et al. (2006)
NA
NA
NA
NA
SF – surface flow; HSSF – horizontal subsurface flow; VSSF – vertical subsurface flow.
rooting zone and contribute to increased N2 O and CO2 production and CH4 consumption (Brix, 1997). Vascular plants
can exchange GHG between the rooting zone and atmosphere (Yavitt and Knapp, 1998). Vegetation and its composition affect the nutrient dynamics and the production, consumption, and transport of GHG and hence their exchange
between wetlands and atmosphere (Ström et al., 2003, 2005;
Søvic et al., 2006; Johansson et al., 2003). They can also affect the biogeochemistry of CWs due to the differences in
their growth and development, longevity, root systems, root
density, root depth, and microbial ecology in the rhizosphere.
As some plant litter decomposes, organic matter with lignocellulose and humic compounds may be released that are
more or less labile or stable in nature than others. Release of
low molecular weight organic matter that is labile in nature
is more likely to produce GHGs than stable forms. For example, Z. latifolia showed higher nutrient removal and CH4
fluxes than P. australis (Inamori et al., 2007). The Z. lotifolia root system is shallow and the activity of methanotrophs
is primarily confined to the top soil. The root systems of P.
australis are deeper, which is more favourable for the oxidization of CH4 .
www.hydrol-earth-syst-sci.net/20/1/2016/
A fluctuating water table in CWs has significant impacts
on GHG dynamics. Pulsing hydrologic regimes decreases
CH4 but increases N2 O emissions (Mander et al., 2011). In
aerobic and anaerobic conditions caused by pulsing hydrology, incomplete nitrification and denitrification increase N2 O
emissions (Healy et al. 2007). However, the effects of pulsing
hydrologic regimes on GHG emissions are contradictory. For
example, intermittent hydrologic regimes decrease both N2 O
(Sha et al., 2011) and CH4 emissions (Song et al., 2010).
Highly contrasting results on gas emissions with fluctuating
water levels have been reported and the controlling mechanisms are unclear (Elberling et al., 2011).
Therefore, the assessment of GHG emissions in various
types of CWs (surface flow, subsurface flow, vertical and
horizontal), vegetation cover (vegetated, non-vegetated) and
species type, and management system employed (HLR, HRT,
soil used, and water table), is necessary in light of the national and global GHG budgets. In addition, such measurements will help scientists, environmental managers, and policy makers adopt environmentally friendly construction and
management of CWs. The enhanced reduction of N2 O to N2
needs further elucidation.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
8
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
Table 4. Carbon dioxide (CO2 ; mg C m−2 d−1 ), CH4 (mg C m−2 d−1 ), and CH4 emission factor (CH4 –C / TOC input × 100) in various
types of CWs; mean standard error ( ± SE) was presented for CH4 emission factor; NA – data not available.
CWs type
Treatment
CO2 emissions
(mg C m−2 d−1 )
CH4 emissions
(mg C m−2 d−1 )
CH4 / TC (%)
References
SF
SF
SF
SF
SF
SF_Finland
SF_Finland
SF_Norway
All SF
Municipal
Domestic
Domestic
Agril. drainage
Dairy wash out
Municipal
Agril runoff
Municipal
NA
0.19
1.13
NA
4250 ± 550
1200 ± 420
3200 ± 560
1400 ± 250
1675 ± 703
5.4
NA
NA
0.88
223 ± 35
29 ± 6.4
350 ± 180
72 ± 28
113 ± 58
NA
26
16
31
9.45
19 ± 4.3
11 ± 5.5
4.8 ± 2.2
16.8 ± 3.8
Tai et al. (2002)
Gui et al. (2007)
Liu et al. (2009)
Wild et al. (2002)
Van der Zaag et al. (2010)
Søvik et al. (2006)
Søvik et al. (2006)
Søvik et al. (2006)
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF
HSSF_Estonia
HSSF_Norway
All HSSF
Domestic
Domestic
Domestic
Domestic
Domestic
Dairy wash out
Domestic
Domestic
Municipal
Municipal
NA
2.54–5.83
5.33
NA
NA
3475 ± 375
0.6–1.7
600
3800 ± 210
790 ± 170
1010 ± 672
1.7–528
0.03–0.40
0.001
0.03
0.29
118 ± 9.0
1.4–4.1
0.48
340 ± 240
130 ± 43
112 ± 74
NA
NA
0.03
4.3
4.0
4.4
0.12–0.23
0.02
NA
9.5 ± 3.3
3.23 ± 1.4
Mander et al. (2005)
Teiter and Mander (2005)
Garcia et al. (2007)
Gui et al. (2007)
Liu et al. (2009)
Van der Zaag et al. (2010)
Søvik et al. (2006)
Mander et al. (2011)
Søvik et al. (2006)
Søvik et al. (2006)
VSSF
VSSF
VSSF
VSSF
VSSF
VSSF
VSSF_Estonia
VSSF_Norway
All VSSF
Domestic
Domestic
Domestic
Domestic
Municipal
Domestic
Municipal
Municipal
5.83–12.13
NA
NA
NA
2662 ± 175
1080
8400 ± 2100
22000 ± 5000
6616 ± 3779
0.60–5.70
16.4
0.013
0.13
33.5 ± 3.2
3.36
110 ± 35
140 ± 160
42.9 ± 23.7
NA
1.68
1.73
NA
0.05
NA
0.39 ± 0.27
1.27 ± 0.31
Teiter and Mander (2005)
Mander et al. (2005)
Gui et al. (2007)
Liu et al. (2009)
Mander et al. (2008)
Mander et al. (2011)
Søvik et al. (2006)
Søvik et al. (2006)
SF – surface flow; HSSF – horizontal subsurface flow; VSSF – vertical subsurface flow.
5
Surface emissions vs. subsurface export of C and N
Dissolved GHG produced in soils and subsoils can be emitted to the atmosphere by transpiration of vascular plants
(from within the rooting zone), ebullition, and diffusion from
soils. Elberling et al. (2011) reported that in wetlands, the
transport of gases through subsoil occurs both via diffusive
transport in the pores and through the vascular plants. Surface emissions of GHG from CWs have recently been recognized and have been commonly measured by chamber methods (Mander et al., 2008, 2011). As is the case with other dissolved pollutants (Dzakpasu et al., 2014), the GHG produced
in CWs can also be transported to the groundwater with the
percolating water and emitted to the atmosphere upon discharge to surface waters (Riya et al., 2010). It can also flow
towards surface waters by advective transport and/or by dispersion of groundwater. Dissolved nutrients can be preferentially leached down into deeper soil layers and groundwater
via different pathways (e.g. root channels). The Nr delivered
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
to groundwater can be transformed in situ to other reactive
or benign forms. Hence, quantification of such Nr loadings
to groundwater and their in situ consumption (e.g. N2 O to
N2 or CH4 to CO2 ) is necessary to understand their environ+
mental consequences. In addition, DON, NO−
3 , NH4 , and
DOC delivered to surface waters can undergo biochemical
reactions and produce N2 O, CO2 , and CH4 in streams and
estuaries. Ström et al. (2007) measured a considerable quantity of CH4 in porewater and found a correlation between
the surface emissions and porewater CH4 concentrations in
vegetated wetlands. Measuring only the surface emissions of
GHG can omit substantial quantities of GHG released from
CWs. For example, Riya et al. (2010) measured emissions
of CH4 and N2 O, accounting for 2.9 and 87 % of the total
emissions. Measuring porewater GHG and linking these to
the surface emissions and subsurface export to groundwater
below CWs will help to estimate a better GHG balance from
both a national and global context. Elberling et al. (2011)
www.hydrol-earth-syst-sci.net/20/1/2016/
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
9
Table 5. Nitrous oxide (N2 O; mg N m−2 d−1 ), CO2 , and CH4 emissions (mg C m−2 d−1 ) in various type of CWs under different plant
types; NA – data not available.
CW type
Wastewater type
Plant type
N2 O
(mg N m−2 d−1 )
CH4
(mg C m−2 d−1 )
HSF
Secondary treated
No plant
3.79 ± 2.64
163 ± 209
municipal
Typha lotifolia
Phalaris arundinacea
Glyceria maxima
Lemna minor
Spirogyra sp.
No plant
Typha atifolia
Phragmites australis
Juncus effusus
No plant
Phragmites
Typha
Phalaris
Phragmites australis
Phragmites australis
2.64 ± 4.09
3.79 ± 3.44
0.76 ± 1.01
1.45 ± 1.18
0.98 ± 1.25
−0.26 ± 2.53
4.94 ± 2.00
7.80 ± 2.53
3.87 ± 1.86
0.04 ± 0.02
0.06 ± 0.03
0.03 ± 0.01
0.01 ± 0.01
15 ± 3.9
264
109 ± 185
212 ± 151
112 ± 178
450 ± 182
107 ± 135
−4.76 ± 61.8
225 ± 47.7
333 ± 76.6
489 ± 46.3
87 ± 6.3
50 ± 7.5
28 ± 3.0
45 ± 6.0
110 ± 35
384
HSF
Sewage treatment
water
HSSF
Domestic
VSSF
VSSF
Municipal
Municipal
CO2
(mg m−2 d−1 )
Reference
Johansson et al. (2003);
Johansson et al. (2004)
NA
NA
NA
NA
NA
4.32 ± 0.73
25.3 ± 4.08
25.1 ± 4.74
26.1 ± 3.00
80 ± 6.3
200 ± 35
235 ± 32
195 ± 31
8400 ± 2100
Ström et al. (2007)
Maltais-Landry et al. (2009)
Søvik et al. (2006)
Mander et al. (2005)
SF – surface flow; HSSF – horizontal subsurface flow; VSSF – vertical subsurface flow.
linked subsurface gas concentrations in wetlands to the surface fluxes using a diffusion model. This demonstrates the
need for future studies on subsurface GHG production, consumption and net GHG emissions in CWs within a climate
change context.
It is important to characterize soils and subsoils’ physical
(e.g. texture, bulk density) and hydraulic (development of a
soil water characteristic curve) properties and to assess their
potential to percolate dissolved nutrients and gases in the
solute phase to the underlying groundwater. To our knowledge, the indirect pathway of GHG emissions from CWs has
never been reported, despite the fact that this would appear
to have a high biogeochemical potential to produce and exchange GHG. The balance between N and C input and output
flows between CWs and aquatic and atmospheric environments, together with the direct and indirect emissions of C
and N species, could be an important input to global C and N
budgets.
6 Hydrogeochemistry below CWs
Constructed wetlands can be designed with or without a clay
liner or a compacted soil bed at the base, which can lead
to large differences in permeability of the underlying layers.
The variation in permeability of a CW soil bed will affect
solute, nutrient, and GHG flows, and their interactions with
the underlying groundwater (Dzakpasu et al., 2012, 2014).
Groundwater hydrogeochemistry below CWs can therefore
provide a unique insight into such interactions. An example of such interactions would be between nutrient-rich water
discharging from CW cells mixing with laterally moving regional groundwater. It should be noted that groundwater can
also discharge into CWs depending on the hydraulic gradiwww.hydrol-earth-syst-sci.net/20/1/2016/
ents. This means that fully screened, multi-level piezometers or boreholes should be installed at such sites to elucidate
groundwater flow direction, hydraulic gradients, and conductivities. Such monitoring networks allow water samples to be
collected and the sources of nutrients in groundwater bodies
below CWs to be identified. The local site hydrology (precipitation, groundwater table fluctuations, and evapotranspiration) has a large impact on the pollutant removal. Hydrogeochemical studies at an accurate spatial and temporal resolution should explain the effects of precipitation on nutrient
removal by dilution as well in situ nutrient turnover. Effective
CW management requires an understanding of the effects of
wetland hydrology on the physical and biochemical attenuation of nutrients in order to assess their impacts on the surface emissions and subsurface export of nutrients and GHG.
Data on the species of N in groundwater below the CWs
are required to provide an in-depth understanding of wetland
ecosystem services, particularly if CWs have the potential
to leak pollutants down into the groundwater (Dzakpasu et
al., 2014). Higher NH+
4 concentrations in groundwater below
the CW than the effluent are often reported (Harrington et
al., 2007; Dzakpasu et al., 2012). Therefore, questions arise
with respect to NH+
4 concentrations in groundwater below
the CWs if they have been transported from CWs. Linking
geochemistry of groundwater below CWs to site hydrology,
water table fluctuations, and soil/subsoil physico-chemical
properties is required to elucidate the major environmental
drivers of C and N removal, and/or pollution swapping. The
quality of groundwater underlying CWs with regards to the
Nr species is largely unknown.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
10
7
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
Methodological developments
To improve the ecosystem services and to minimize the pollution swapping of CWs, quantification of N cycling is crucial. Measurement of GHG using the closed chamber method
is widely used, but has large uncertainty in estimating the diurnal variability due to internal changes in temperature and
physical access to the chambers over a 24 h time period. Gas
ebullition and diffusion measurements are quite challenging
in CWs covered by vegetation, because of the difficulties in
estimation of gas transfer velocity. Application of the eddycovariance method is not appropriate for most CWs, as it requires a large surface area (> several ha) to avoid contribution
of surrounding area and complication of GHG foot printing.
A combination of chamber, ebullition, and diffusion methods
in a single system could minimize the uncertainly in GHG estimation. The methane ebullition measurement was found to
be similar to surface emissions by the chamber method, but
N2 O and CO2 ebullition measurements were lower than the
surface emissions (Søvik et al., 2006).
The use of in situ microcosm studies and soil core incubation methods may give a better estimation of N2 O, CO2 ,
and CH4 production and consumption than existing methods. With the recent advancement of isotope pairing and dilution techniques, single or simultaneously occurring C and
N transformation processes can be quantified in laboratory
or in situ conditions (Huygens et al., 2013; Müller et al.,
2014). The isotope technique relies on the introduction of
a known amount of 14 C and or 15 N into the CW and then
quantification of C and N concentrations and isotopic compositions through different C and N pools after incubation
for a specific period. Laboratory methods involve collection
of intact soil/sediment cores, with subsequent incubation in
the laboratory. In situ field techniques involve the release
of a 14 C / 15 N solution in the CW soils. Incubation of intact soil cores with differentially labelled 15 NH14
4 NO3 and
14 NH15 NO can be used to quantify the rates of different
3
4
N transformation processes (Rütting and Müller, 2008). The
quantification of simultaneously occurring N transformation
rates rely on the analysis with appropriate 15 N-tracing models. In recent years, 15 N-tracing techniques have evolved, and
are now able to identify process-specific NO−
2 pools (Rütting and Müller 2008), pathway-specific N2 O production,
and emission, as well as N2 O : N2 ratios (Müller et al., 2014).
Traditional techniques for investigation of gross N dynamics in sediments (Blackburn, 1979) may be combined with
the latest 15 N-tracing techniques, where all N transformation rates are included (Huygens et al., 2013). Thus, current
models should consider processes such as anammox and/or
deamox, and then be tested in CWs under various operational
conditions. Denitrification in porewater samples can be
measured by analysing samples for dissolved N2 in a membrane inlet mass spectrometer (MIMS; Kana et al., 1994)
and N2 O in a gas chromatograph (GC; Jahangir et al., 2012).
The studies of natural abundance of 15 N and 18 O (δ 15 N and
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
δ 18 O) in NO−
3 is an insightful tool for the investigation of the
sources, fate, and transformational processes of N in a system
(e.g. in shallow groundwater; Baily et al., 2011). The in situ
NO−
3 push–pull method has been used to determine denitrification in shallow groundwater (< 3 m) in riparian wetlands
(Addy et al., 2002; Kellogg et al., 2005) and in deep groundwater in arable/grassland (Jahangir et al., 2013).
Isotope-based techniques can also be extended to other
elements; e.g., a 33 P-tracing model has been developed recently to study phosphorus (P) cycle in soil (Müller and
Bünemann, 2014). These techniques can be applied in the
study of C, N, and P biogeochemistry in aquatic environments. In addition, measurements of DOC and gases (CO2
and CH4 ) will provide insights into the C consumption and
transformation associated with the N transformations. Carbon and N dynamics are influenced by the interacting effects of soil conditions with microbial community structure
and functioning. Microbial functioning involves transcription of genes, translation of messenger RNA, and activity of
enzymes (Firestone et al., 2012). As such, activities of microbial communities under various environmental conditions
and how these contribute to C and N dynamics is a very important area of future research (Müller and Clough, 2014).
Molecular approaches can be important tools for identifying
and quantifying the genes that code for enzyme-mediating
C and N cycles (Peterson et al., 2012). These tools help
assess the relationships among genes, environmental controllers, and the rates of C and N processes. The scientific
tools and multi-disciplinary techniques are now available to
better understand C and N transformation rates, processes,
and factors controlling the unwanted emission of N and C
products to the environment.
8
Conclusions and recommendations
The transformational processes on a mixture of contaminants within and below CWs can cause pollution swapping.
A holistic assessment of C and N dynamics in CWs is needed
to fully understand their removal, transport, and impact on
water quality and emissions to atmosphere. Mixed contaminants entering CWs and those formed within and underneath
CWs during transformational processes must be considered
in future studies. The overall balance of these constituents
will determine whether a CW is a pollution source or a sink.
This will necessitate a higher degree of multi-level spatial
and temporal monitoring and the use of multi-disciplinary
in and ex situ techniques to fully characterize all pathways
of C and N loss. At this time we cannot suggest any design
optima in terms of nutrient removal and GHG mitigation because empirical information is not yet abundant. To do this,
transformation kinetics of C and N and net GHG emissions
through all possible pathways are required to provide a holistic assessment. However, a combination of various types of
CW and plant types could provide higher removals and lower
www.hydrol-earth-syst-sci.net/20/1/2016/
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
11
Figure 1. Conceptual model showing the current state of knowledge of C and N dynamics in constructed wetlands treating wastewater and
the specific experimental work that needs to be undertaken in the future; SF – surface flow; HSSF – horizontal subsurface flow; VSSF –
vertical subsurface flow; HLR – hydraulic loading rate; HTR – hydraulic retention time; ? – not known or very little known.
GHG emissions. A conceptual model highlighting the current state of knowledge in this area and the research gaps is
presented in Fig. 1.
Subsurface export of nutrients and GHG to groundwater
should be accounted for in CW management. Reducing the
saturated hydraulic conductivity below the wetland bed will
help reduce nutrients leaching to groundwater. The reactive
versus the benign forms of the N transformation products
should be evaluated. Data on when, where, and the rates at
which denitrification, deamox, and anammox occur in CWs
are needed, as well as identification of the key factors that
control such processes. The provenance of NH+
4 in groundwater below CW cells and its impact on down-gradient receptors needs further elucidation. Constructed wetlands have
the potential to produce N2 O, DON, DOC, dissolved inorganic C (DIC), CO2 , and CH4 , which may be exported to
fresh waters via groundwater and degassed upon discharge
to surface waters. Moreover, the DOC and DIC transferred
to the fresh water sediments (rivers and lakes) can produce
GHG that, in turn, emit to the atmosphere. The amount of C
and N exported from terrestrial ecosystems via the subsurface pathway to fresh waters has been the missing piece of
our understanding of global C and N budgets. It is clear that
data on the various C and N species, along with the GHG
emissions, are crucial to make a robust input–output balance
of C and N in CWs. Spatial and temporal variations of GHG
emissions in CWs under different management systems are
also critical to get much more rigorous estimates of emission
factors. These data will reduce the existing uncertainties in
global C and N budgets.
www.hydrol-earth-syst-sci.net/20/1/2016/
Managing wetting and drying spells (pulsing hydrology)
in CWs can enhance NH+
4 removal. Similarly, oxidation of
organic C will increase CO2 production and, in anaerobic
conditions, may be reduced to CH4 . This requires more research into the C and N cycle processes over the wetting and
drying spells, which is now possible with the advancement
in 14 C / 15 N-tracing and modelling techniques. The selection
of appropriate plant species is important to optimize nutrient removal, sequester C, and decrease GHG emissions, but
more research is needed across species and geographical locations. Further research is also needed to investigate the impacts of hydraulic retention time on nutrient dynamics. Rates
of nutrient accumulation or fixation in soils and their in situ
transformation in CWs need to be quantified to evaluate their
contribution to C sequestration and GHG emissions.
Author contributions. The first author, M. M. R. Jahangir has reviewed articles in the relevant area, analysed results, identified
knowledge gaps, and prepared the draft paper. All co-authors were
directly involved in preparation of the paper and edited the paper
for its improvement.
Acknowledgements. The research was funded by Irish Research
Council and Department of Agriculture, Food and Marine in
Association with The University of Dublin, Trinity College.
Edited by: M. Hipsey
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
12
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
References
á Norði, K. and Thamdrup, B.: Nitrate-dependent anaerobic
methane oxidation in a freshwater sediment, GeochiM. Cosmochim. Acta, 132, 141–150, 2014.
Addy, K., Kellogg, D. Q., Gold, A. J., Groffman, P. M., Ferendo, G.,
and Sawyer, C.: In situ push-pull method to determine groundwater denitrification in riparian zones, J. Environ. Qual., 31, 1017–
1024, 2002.
Bachand, P. A. M. and Horne, A. J.: Denitrification in constructed
free-water surface wetlands: II. Effects of vegetation and temperature, Ecol. Eng., 14, 17–32, 2000.
Baily, A., Rock, L., Watson, C. J., and Fenton, O.: Spatial and temporal variations in groundwater nitrate at an intensive dairy farm
in south-east Ireland: Insights from stable isotope data, Agril.
Ecosysts. Environ., 308–318, 2011.
Beaulieu, J. J., Tank, J. L., Hamilton, S. K., Wollheim, W. M., Hall,
R. O., and Mulholland, P. J.: Nitrous oxide emission from denitrification in stream and river networks, P. Natl. Acad. Sci., 108,
214–219, 2011.
Blackburn, T. H.: Methods for measuring rates of NH+
4 turnover in
anoxic marine sediments, using a 15 N-NH+
dilution
technique,
4
Appl. Environ. Microbiol., 37, 760–765, 1979.
Bojcevska, H. and Tonderski, K.: Impact of loads, season, and plant
species on the performance of a tropical constructed wetland polishing effluent from sugar factory stabilization ponds, Ecol. Eng.,
29, 66–76, 2007.
Brix, H.: Do macrophytes play a role in constructed treatment wetlands, Water Sci. Technol., 35, 11–17, 1997.
Brix, H., Sorrell, B. K., and Lorenzen, B.: Are Phragmitesdominated wetlands a net source or net sink of greenhouse
gases?, Aquat. Bot., 69, 313–324, 2001.
Burgin, A. J., Hamilton, S. K., Gardner, W. S., and McCarthy, M.
J.: Nitrate reduction, denitrification, and dissimilatory nitrate reduction to ammonium in wetland sediments, in: Methods in Biogeochemistry of Wetlands, edited by: DeLaune, R. D., Reddy, K.
R., Richardson, C. J., and Megonigal, J. P., SSSA Book Series,
no. 10, Madison, USA, 307–325, 2013.
Calheiros, C. S. C., Rangel, A. O. S. S., and Castro, P. M. L.: Constructed wetland systems vegetated with different plants applied
to the treatment of tannery wastewater, Water Res., 41, 1790–
1798, 2007.
Clair, T. A., Arp, P., Moore, T. R., Dalva, M., and Meng, F. R.:
Gaseous carbon dioxide and methane, as well as dissolved organic carbon losses from a small temperate wetland under a
changing climate, Environ. Pollut., 116, S143–S148, 2002.
Conley, D. J., Paerl, H. W., Howarth, R. W. Boesch, D. F.,
Seitzinger, S. P., Havens, K. E., Lancelot, C., and Likens, G. E.:
Controlling eutrophication: Nitrogen and phosphorus, Science,
323, 1014–1015, 2009.
da Motta Marques D. M. L., Leite, G. R., and Giovannini, S. G.
T.: Performance of two macrophyte species in experimental wetlands receiving variable loads of anaerobically treated municipal
wastewater, Water Sci. Technol., 44, 311–316, 2001.
Drizo, A., Frost, C. A., Grace, J., and Smith, K. A.: Physicochemical screening of phosphate-removing substrates for use in
constructed wetland systems, Wat. Res., 33, 3595–3602, 1999.
Dzakpasu, M., Hofmann, O., Scholz, M., Harrington, R., Jordan,
S. N., and McCarthy, V.: Nitrogen removal in an integrated con-
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
structed wetland treating domestic wastewater, J. Environ. Sci.
Health, Part A, 46, 742–750, 2011.
Dzakpasua, M., Scholz, M., Harrington, R., Jordan, S. N., and McCarthy, V.: Characterising infiltration and contaminant migration beneath earthen-lined integrated constructed wetlands, Ecol.
Eng., 41, 41–51, 2012.
Dzakpasu, M., Scholz, M., Harrington, R., McCarthy, V., and Jordan, S.: Groundwater Quality Impacts from a Full-Scale Integrated Constructed Wetland, Groundw. Monit. Remote, 34, 51–
64, 2014.
Elberling, B., Louise A., Christian, J. J., Hans, P. J., Michael K.,
Ronnie N. G., and Frants, R. L.: Linking soil O2 , CO2 , and CH4
concentrations in a wetland soil: implications for CO2 and CH4
fluxes, Environ. Sci. Technol., 45, 3393–3399, 2011.
Fazzolari, C. E., Nicolardot, B., and Germon, J. C.: Simultaneous
effects of increasing levels of glucose and oxygen partial pressures on denitrification and dissimilatory reduction to ammonium in repacked soil cores, Eur. J. Soil Biol., 34, 47–52, 1998.
Fenton, O., Healy, M. G., Brennan, F., Jahangir, M. M. R., Lanigan, G. J., Richards, K. G., Thornton, S. F., and Ibrahim, T. G.:
Permeable reactive interceptors – blocking diffuse nutrient and
greenhouse gas losses in key areas of the farming landscape, J.
Agric. Sci., 152, S71–S81, 2014.
Fey, A., Benckiser, G., and Ottow, J. C. G.: Emissions of nitrous oxide from a constructed wetland using a ground filter and macrophytes in waste-water purification of a dairy farm, Biol. Fertil.
Soils, 29, 354–359, 1999.
Findlay, S., Groffman, P., and Dye, S.: Effects of Phragmites australis removal on marsh nutrient cycling, Wetlands Ecol. Manage., 11, 157–165, 2003.
Firestone, M., Blazewicz, S., Peterson, D. G., and Placella, S.: Can
molecular microbial ecology provide new understanding of soil
nitrogen dynamics, edited by: Richards, K. G., Fenton, O., Watson, C. J., in: Proceedings of the 17th Nitrogen Workshop, 11–
14, 2012.
Galloway, J. N., Aber, J. D., Erisman, J. W., Seitzinger, S. P.,
Howarth, R. W., Cowling, E. B., and Cosby, B. J.: The nitrogen
cascade, BioSci., 53, 341–356, 2003.
Garcia, J., Capel, V., Castro, A., Ruiz, I., and Soto, M.: Anaerobic biodegradation tests and gas emissions from subsurface flow
constructed wetlands, Bioresour. Technol., 98, 3044–3052, 2007.
Gray, N. F.: Drinking water quality: problems and solutions, 2nd
edn., Cambridge Univ. Press, Cambridge, UK, 116–134, 2008
Grünfeld, S. and Brix, H.: Methanogenesis and methane emissions:
effects of water table, substrate type and presence of Phragmites
australis, Aquat,. Bot., 64, 63–75, 1999.
Gui, P., Inamori, R., Matsumura, M., and Inamori, Y.: Evaluation
of constructedwetlands by waste water purification ability and
greenhouse gas emissions, Water Sci. Technol., 56, 49–55, 2007.
Haroon, M. F., Hu, S., Shi, Y., Imelfort, M., Keller, J., Hugenholtz,
P., Yuan, Z., and Tyson, G. W.: Anaerobic oxidation of methane
coupled to nitrate reduction in a novel archaeal lineage, Nature,
500, 567—570, 2013.
Harrington, R., Carroll, P., Carty, A. H., Keohane, J., and Ryder, C.:
Integrated constructed wetlands: concept, design, site evaluation
and performance, Int. J. Water, 3, 243–255, 2007.
Healy, M. G., Rodgers, M., and Mulqueen, J.: Treatment of dairy
wastewater using constructed wetlands and intermittent sand filters, Bioresour. Technol., 98, 2268–2281, 2007.
www.hydrol-earth-syst-sci.net/20/1/2016/
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
Healy, M. G., Ibrahim, T. G., Lanigan, G., Serrenho, A. J., and Fenton, O.: Nitrate removal rate, efficiency and pollution swapping
potential of different organic carbon media in laboratory denitrification bioreactors, Ecol. Eng., 40, 198–209, 2011.
Healy, M. G., Barrett, M., Lanigan, G. J., João Serrenho, A.,
Ibrahim, T. G., Thornton, S. F., Rolfe, S. A., Huang, W. E., and
Fenton, O.: Optimizing nitrate removal and evaluating pollution
swapping trade-offs from laboratory denitrification bioreactors,
Ecol. Eng., 74, 290–301, 2014
Huygens, D., Trimmer, M., Rütting, T., Müller, C., Heppell, C. M.,
Lansdown, K., and Boeckx, P.: Biogeochemical N cycling in wetland ecosystems: 15 N isotope techniques, Methods in biogeochemistry of wetlands, edited by: Reddy, K. R., Megonigal, J.
P., and Delaune, R. D., Soil Sci. Soc. Amer., 30, 553–591, 2013.
Inamori, R., Gui, P., Dass, P., Matsumura, M., Xu, K.Q., Kondo,
K., Ebie, Y., and Inamori, Y.: Investigating CH4 and N2 O emissions from eco-engineering wastewater treatment processes using constructed wetland microcosms, Proc. Biochem, 42, 363–
373, 2007.
IPCC: 2013 Supplement to the 2006 IPCC guidelines for national
greenhouse gas inventories: wetlands, edited by: Hirashi, T.,
Krug, T., Tanabe, K., Srivastava, N., Baasansuren, J., Fukuda,
M., Troxler, T. G., Switzerland, IPCC Task Force on National
Greenhouse Gas Inventories, 354, 2013.
Islas-Lima, S., Thalasso, F., and Gomez-Hernandez, J.: Evidence of
anoxic methane oxidation coupled to denitrification, Water Res.,
38, 13–16, 2004.
Jahangir, M. M. R., Johnston, P., Grant, J., Somers, C., Khalil, M. I.,
and Richards, K. G.: Evaluation of headspace equilibration methods for measuring greenhouse gases in groundwater, J. Environ.
Manage., 111, 208–212, 2012.
Jahangir, M. M. R., Johnston, P., Addy, K., Khalil, M. I., Groffman,
P., and Richards, K. G.: Quantification of in situ denitrification
rates in groundwater below an arable and a grassland system, Water Air Soil Pollut., 224, 1693, doi:10.1007/s11270-013-1693-z,
2013.
Johansson, A. E., Kasimir-Klemedtsson, A., Klemedtsson, L., and
Svensson, B. H.: Nitrous oxide exchanges with the atmosphere of
a constructed wetland treating wastewater, Tellus, 55B, 737–750,
2003.
Johansson, A. E., Gustavsson, A. M., Öquist, M. G., and Svensson,
B. H.: Methane emissions from a constructed wetland treating
wastewater – seasonal and spatial distribution and dependence
on edaphic factors, Water Res., 38, 3960–3970, 2004.
Johansson, E.: Constructed wetlands and deconstructed discorses–
greenhouse gas fluxes and discorses on purifying capacities,
Dept. Water Environ. Studies, Linköping Univ., Sweden, 2002.
Kalyuzhnyi, S., Gladchenko, M., Mulder, A., and Versprille, B.:
DEAMOX-new biological nitrogen, Water Res., 40, 3637–3645,
2006.
Kana, T. M., Darkangelo, C., Hunt, M. D., Oldham, J. B., Bennett,
G. E, and Cornwell, J. C.: Membrane inlet mass spectrometer
for rapid high precision determination N2 , O2 and Ar in environmental water samples, Anal. Chem., 66, 4166–4170, 1994.
Kato, K., Koba, T., Ietsugu, H., Saigusa, T., Nozoe, T., Kobayashi,
S., Kitagawa, K., and Yanagiya, S.: Early performance of hybrid
reed bed system to treat milking parlour wastewater in cold climate in Japan, in: 10th international conference wetland systems
www.hydrol-earth-syst-sci.net/20/1/2016/
13
for water pollution control, MAOTDR, Lisbon, Portugal, 1111–
1118, 2006.
Kellogg, D. Q., Gold, A. J., Groffman, P. M., Addy, K., Stolt, M. H.,
and Blazejewski, G.: In situ groundwater denitrification in stratified, permeable soils underlying riparian wetlands, J. Environ.
Qual., 34, 524–533, 2005.
Kelso, B. H. L., Smith, R. V., Laughlin, R. J., and Lennox, S. D.:
Dissmilatory nitrate reduction in anaerobic sediments leading to
river nitrate accumulation, Appl. Enviro. Microbiol., 63, 4679–
4685, 1997.
Korom, S. F.: Natural denitrification in the saturated zone: a review,
Water Resour. Res., 28, 1657–1668, 2002.
Koskiaho, J., Ekholm, P., Räty, M., Riihimäki, J., and Puustinen,
M.: Retaining agricultural nutrients in constructed wetlands- experiences under boreal conditions, Ecol. Eng., 20, 89–103, 2003.
Laanbroek, H. J.: Bacterial cycling of minerals that affect plant
growth in waterlogged soils: a review, Aquat. Bot., 38, 109–125,
1990.
Laanbroek, H. J.: Methane emission from natural wetlands: interplay between emergent macrophytes and soil microbial processes, A mini-review, Ann. Bot., 105, 141–153, 2010.
Langergraber, G.: Modeling of processes in subsurface flow constructed wetlands: A review. Vadose Zone, J. 7, 830-842, 2008.
Lee, C., Fletcher, T. D., and Sun, G.: Nitrogen removal in constructed wetland systems, Eng. Life Sci., 9, 11–22, 2009.
Liikanen, A., Huttunen, J. T., Kaijalainen, S. M., Heikkinen, K.
Vdisdinen, T. S., Nykiinen, H., and Martikainen, P. J.: Temporal
and seasonal changes in greenhouse gas emissions from a constructed wetland purifying peat mining runoff water, Ecol. Eng.,
26, 241–251, 2006.
Liu, C., Xu, K., Inamori, R., Ebie, Y., Liao, J., and Inamori, Y.:
Pilot-scale studies of domestic wastewater treatment by typical
constructed wetlands and their greenhouse gas emissions, Front.
Environ. Sci. Eng. China, 3, 477–482, 2009.
Luederitz, V., Eckert, E., Lange-Weber, M., Lange, A., and Gersberg, R. M.: Nutrient removal efficiency and resource economics
of vertical flow and horizontal flow constructed wetlands, Ecol.
Eng., 18, 157–171, 2001.
Luo, W. G., Wang, S. H., Huang, J., and Qian, W. Y.: Denitrification by using subsurface constructed wetland in low temperature,
China Water Wastewater, 21, 37–40, 2005.
Makino, W., Cotner, J. B., Sterner, R. W., and Elser, J. J.: Are bacteria more like plants or animals? Growth rate and resource dependence of bacterial C:N:P stoichiometry, Funct. Ecol., 17, 121–
130, 2003.
Maltais-Landry, G., Maranger, R., Brisson, J., and Chazarenc, F.:
Greenhouse gas production and efficiency of planted and artificially aerated constructed wetlands, Environ. Pollut., 157, 748–
754, 2009.
Mander, Ü., Lõhmus, K., Teiter, S., Nurk, K., Mauring, T., and
Augustin, J.: Gaseous fluxes from subsurface flow constructed
wetlands for wastewater treatment, J. Environ. Sci. Health A 40,
1215–1226, 2005.
Mander, U., Lõhmus, K., Teiter, S., Mauring, T., Nurk, K., and Augustin, J.: Gaseous fluxes in the nitrogen and carbon budgets of
subsurface flow constructed wetlands, Sci. Total Environ., 404,
343–353, 2008.
Mander, Ü., Maddison, M., Soosaar, K., and Karabelnik, K.: The
impact of intermit-tent hydrology and fluctuating water table on
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
14
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
greenhouse gas emissions from subsurface flow constructed wetlands for wastewater treatment, Wetlands, 31, 1023–1032, 2011.
Mander, Ü., Dotro, G., Ebie, Y., Towprayoon, S., Chiemchaisri, C.,
Nogueira, S. F., Jamsranjav, B., Kasak, K, Truu, J., Tournebize,
J., and Mitsch, W. J.: Greenhouse gas emission in constructed
wetlands for wastewatertreatment: A review, Ecol. Eng., 66, 19–
35, 2014.
Minett, D. A., Cook, P. L. M., Kessler, A. J., and Cavagnaro, T.
R.: Root effects on the spatial and temporal dynamics of oxygen
in sand-based laboratory-scale constructed biofilters, Ecol. Eng.,
58, 414–422, 2013.
Mitsch, W. J. and Gosselink, J. G.: Wetlands, 3rd Edn. New York,
John Wiley & Sons, p. 936, 2000.
Müller, C. and Bünemann, E. K.: A 33 P tracing model for quantifying gross P transformation rates in soil, Soil Biol. Biochem., 76,
218–226, 2014.
Müller, C. and Clough, T. J.: Advances in understanding nitrogen
flows and transformations: gaps and research pathways, J. Agril.
Sci., 152, S1, 34–44, 2014.
Müller, C., Laughlin, R. J., Spott, O., and Rütting, T.: Quanti?cation
of N2 O emission pathways via a 15 N tracing model, Soil Biol.
Biochem., 72, 44–54, 2014.
Mustafa, A. and Scholz, M.: Nutrient accumulation in Typha latifolia L. and sediment of a representative integrated constructed
wetland, Water Air Soil Pollut., 219, 329–341, 2011.
Nguyen, L. M.: Organic matter composition, microbial biomass
and microbial activity in gravel-bed constructed wetlands treating farm dairy wastewaters, Ecol. Eng., 16, 199–221, 2000.
O’Luanaigh, N. D., Goodhue, R., and Gill, L. W.: Nutrient removal
from on-site domestic wastewater in horizontal subsurface flow
reed beds in Ireland, Ecol. Eng., 36, 1266–1276, 2010.
Obarska-Pempkowiak, H. and Gajewska, M.: The dynamics of processes responsible for transformation of nitrogen compounds in
hybrid wetlands systems in a temperate climate, in: WetlandsNutrients, Metals and Mass Cycling, edited by: Vymazal, J.,
Backhuys Publishers, Leiden, The Netherlands, 129–142, 2003.
Pangala, S. R., Reay, D. S., and Heal, K. V.: Mitigation of methane
emissions from constructed farm wetlands, Chemosphere, 78,
493–499, 2010.
Parkin, T. B.: Soil microsites as a source of denitrification variability, Soil Sci. Am. J., 51, 1194–1199, 1987.
Payne, E. G. I., Fletcher, T. D., Cook, P. L. M., Deletic, A., and Hatt,
B. E.: Processes and drivers of nitrogen removal in stormwater biofiltration, Crit. Rev. Environ. Sci. Technol., 44, 796–846,
2014a.
Payne, E. G. I., Fletcher, T. D., Russell, D. G., Grace, M. R., Cavagnaro, T. R., Evrard, V., Deletic, A., Hatt, B. E., and Cook,
P. L. M.: Temporary storage or permanent removal? The division of nitrogen between biotic assimilation and denitrification in
stormwater biofiltration systems, PLoS ONE, 9, e90890, 2014b.
Peterson, D. G., Blazewicz, S., Herman, D. J., Firestone, M., Turetsky, M., and Waldrop, M.: Abundance of microbial genes associated with nitrogen cycling as indices of biogeochemical process
rates across a vegetation gradient in Alaska, Environ. Microbiol.,
14, 993–1008, 2012.
Pett-Ridge, J., Silver, W. L., and Firestone, M. K.: Redox fluctuations frame microbial community impacts on N-cycling rates in
humid tropical forest soil, Biogeochem., 81, 95–110, 2006.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016
Poach, M. E., Hunt, P. G., Vanotti, M. B., Stone, K. C., Matheny,
T. A., Johnson, M. H., and Sadler, E. J.: Improved nitrogen treatment by constructed wetlands receiving partially nitrified liquid
swine manure, Ecol. Eng., 20, 183–197, 2003.
Rabalais, N. N., Díaz, R. J., Levin, L. A., Turner, R. E., Gilbert, D.,
and Zhang, J.: Dynamics and distribution of natural and humancaused hypoxia, Biogeosciences, 7, 585–619, doi:10.5194/bg-7585-2010, 2010.
Reay, D. S.: Fertilizer ‘solution’ could turn local problem global
– Protecting soil and water from pollution may mean releasing
more greenhouse gas, Nature, 427, 485–485, 2004.
Reddy, K. R. and Delaune, R. D.: Biogeochemistry of Wetlands:
Science and Applications, CRC Press, 800 pp., 2008.
Riya, S., Zhou, S., Nakashima, Y., Terada, A., and Hosomi, M.:
Direct and indirect greenhouse gas emissions from vertical flow
constructed wetland planted with forage rice, Kagaku Kogaku
Ronbunshu, 36, 229–236, 2010.
Rütting, T. and Müller, C.: Process-specific analysis of nitrite dynamics in a permanent grassland soil by using a Monte Carlo
sampling technique, Euro. J. Soil Sci., 59, 208–215, 2008.
Rütting, T., Boeckx, P., Müller, C., and Klemedtsson, L.: Assessment of the importance of dissimilatory nitrate reduction to ammonium for the terrestrial nitrogen cycle, Biogeosci., 8, 1169–
1196, 2011.
Seitzinger, S. P., Sanders, R. W., and Styles, R.: Bioavailability of
DON from natural and anthropogenic sources to estuarine plankton, Limnol. Oceanogr., 47, 353–366, 2002.
Sha, C. Y., Mitsch, W. J., Mander, Ü., Lu, J. J., Batson, J., Zhang,
L., and He, W. S.: Methane emissions from freshwater riverine
wetlands, Ecol. Eng., 37, 16–24, 2011.
Shamir, E., Thompson, T. L., Karpisak, M. M., Freitas, R. J., and
Zauderer, J.: Nitrogen accumulation in a constructed wetland for
dairy wastewater treatment, J. Amer. Water Resour. Assoc., 37,
315–325, 2001.
Song, K., Lee, S. H., Mitsch, W. J., and Kang, H.: Different responses of denitrification rates and denitrifying bacterial communities to hydrological pulses in created wetlands, Soil Biol.
Biochem., 42, 1721–1727, 2010.
Song, K., Lee, S. H., and Kang, H.: Denitrification rates and community structure of denitrifying bacteria in newly constructed
wetland, Eur. J. Soil Biol., 47, 24–29, 2011.
Søvik, A. K., Augustin, J., Heikkinen, K., Huttunen, J. T., Necki, J.
M., Karjalainen, S. M., Klove, B., Liikanen, A., Mander, U., Puustinen, M., Teiter, S., and Wachniew, P.: Emission of the greenhouse gases nitrous oxide and methane from constructed wetlands in Europe. J. Environ, Qual., 35, 2360–2373, 2006.
Stadmark, J. and Leonardson, L.: Emissions of greenhouse gases
from ponds constructed for nitrogen removal, Ecol. Eng., 25,
542–551, 2005.
Ström, L., Ekberg, A., Mastepanov, M., and Christensen, T. R.: The
effect of vascular plants on carbon turnover and methane emissions from a tundra wetland, Global Change Biol. 9, 1185–1192,
2003.
Ström, L., Mastepanov, M., and Christensen, T. R.: Species specific
effects of vascular plants on carbon turnover and methane emissions from wetlands, Biogeochem., 75, 65–82, 2005.
Ström, L., Lamppa, A., and Christensen, T. R.: Greenhouse gas
emissions from a constructed wetland in southern Sweden, Wetlands Ecol. Manage., 15, 43–50, 2007.
www.hydrol-earth-syst-sci.net/20/1/2016/
M. M. R. Jahangir et al.: Carbon and nitrogen dynamics and greenhouse gas emissions in constructed wetlands
Tai, P. D., Li, P. J., Sun, T. H., He, Y. W., Zhou, Q. X., Gong, Z. Q.,
Mizuochi, M., and Inamori, Y.: Greenhouse gas emissions from a
constructed wetland for municipal sewage treatment, J. Environ.
Sci. China, 14, 27–33, 2002.
Takaya, N.: Dissimilatory nitrate reduction metabolisms and their
control in fungi, J. Biosci. Bioeng., 94, 506–510, 2002.
Tanner, C. C., Nguyen, M. L., and Sukias, J. P. S.: Nutrient removal by a constructed wetland treating subsurface drainage
from grazed dairy pasture, Agric. Ecosysts. Environ., 105, 145–
162, 2005.
Teiter, S. and Mander, Ü.: Emission of N2 O, N2 , CH4 and CO2
from constructed wetlands for wastewater treatment and from riparian buffer zones, Ecol. Eng., 25, 528–541, 2005.
Thayalakumaran, T., Bristow, K. L., Charlesworth, P. B., and Fass,
T.: Geochemical conditions in groundwater systems: Implications for the attenuation of agricultural nitrate, Agril. Water Manage., 95, 103–115, 2008.
Toet, S., Richards, S. P., van Logtestijn, Kamp, R., Schreijer, M.,
and Verhoeven, J. T. A.: The effect of hydraulic retention time
on the performance of pollutants from sewage treatment plant
effluent in a surface flow constructed wetland system, Wetlands,
25, 375–391, 2005.
Van der Zaag, A. C., Gordon, R. J., Burton, D. L., Jamieson, R. C.,
and Stratton, G. W.: Greenhouse gas emissions from surface flow
and subsurface flow constructed wetlands treating dairy wastewater, J. Environ. Qual., 39, 460–471, 2010.
Van Oostrom, A. J. and Russell, J. M.: Denitrification in constructed
wastewater wetlands receiving high concentrations of nitrate,
Water Sci. Technol., 29, 7–14, 1994.
Verhoeven, J. T. A. and Meuleman, A. F. M.: Wetlands for wastewater treatment: Opportunities and limitations, Ecol. Eng., 12, 5–
12, 1999.
Vymazal, J.: Removal of nutrients in various types of constructed
wetlands, Sci. Total Environ., 380, 48–65, 2007.
Vymazal, J. and Kröpfelová, L.: Types of constructed wetlands for
wastewater treatment in: Wastewater Treatment in Constructed
Wetlands with Horizontal Sub-Surface Flow, Environmental Pollution Springer, Heidelberg, 121–202, 2010.
Wang, H., Huang, C., Ge, Y., zhi Wu, J., and Chang, J.: The Performance of Species Mixtures in Nitrogen and Phosphorus Removal at Different Hydraulic Retention Times, Pol. J. Environ.
Stud., 23, 917–922, 2014.
Weerakoon, G. M .P. R., Jinadasa, K. B. S. N., Herath, G. B. B.,
Mowjood, M. I. M., and van Bruggen, J. J. A.: Impact of the hydraulic loading rate on pollutants removal in tropical horizontal
subsurface flow constructed wetlands, Ecol. Eng., 61, 154–160,
2013.
www.hydrol-earth-syst-sci.net/20/1/2016/
15
Whalen, S. C.: Biogeochemistry of methane exchange between natural wetlands and the atmosphere, Environ. Eng. Sci., 22, 73–94,
2005.
Wild, U., Lenz, A., Kamp, T., Heinz, S., and Pfadenhauer, J.: Vegetation development, nutrient removal and trace gas fluxes in
constructed Typha wetlands, in: Natural wetlands for wastewater treatment in cold climates, edited by: Mander, Ü., Jenssen P.
D., Adv. Ecol. Sci., Southampton, Boston, WIT Press, 101–126,
2002.
Wunderlin, P., Lehmann, M. F., Siegrist, H., Tuzson, B., Joss, A.,
Emmenegger, L., and Mohn, J.: Isotope signatures of N2 O in a
mixed microbial population system: constraints on N2 O producing pathways in wastewater treatment. Environ. Sci. Technol., 47,
1339–1348, 2013.
Wynn, T. M. and Liehr, S. K.: Development of a constructed
subsurface-flow wetland simulation model, Ecol. Engg., 16,
519–536, 2001.
Xue, Y., Kovacic, D. A., David, M. B., Gentry, L. E., Mulvaney, R.
L., and Lindau, C. W.: In situ measurements of denitrification in
constructed wetlands, J. Environ. Qual., 28, 263–269, 1999.
Yan, C., Zhang, H., Li, B., Wang, D., Zhao, Y., and Zheng, Z.: Effects of influent C/N ratios on CO2 and CH4 emissions from vertical subsurface flow constructed wetlands treating synthetic municipal wastewater, J. Hazardous Mat., 203–204, 188–194, 2012.
Yavitt, J. B. and Knapp, A. K.: Aspects of methane flow from sediment through emergent cattail (Typha latifolia) plants, New Phytol., 139, 495–503, 1998.
Yin, S., Shen, Q., Tang, Y., and Cheng, L.: Reduction of nitrate to
ammonium in selected paddy soils in China, Pedosphere, 8, 221–
228, 1998.
Zhang, J., Shao, W. S., He, M., Hu, H. Y., and Gao, B. Y.: Treatment performance and enhancement of subsurface constructed
wetland treating polluted river water in winter, Environ. Sci., 27,
1560–1564, 2006.
Zhao, Y., Zhang, Y., Ge, Z., Hu, C., and Zhang, H.: Effects of influent C/N ratios on wastewater nutrient removal and simultaneous
greenhouse gas emission from the combinations of vertical subsurface flow constructed wetlands and earthworm eco-filters for
treating synthetic wastewater, Environ. Sci. Proc. Impacts, 16,
567–575, 2014.
Hydrol. Earth Syst. Sci., 20, 1–15, 2016